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1 Katholieke Universiteit Leuven Faculteit Bio-ingenieurswetenschappen ',66(57$7,21(6'($*5,&8/785$ Doctoraatsproefschrift nr. 739 aan de faculteit Bio-ingenieurswetenschappen van de K.U.Leuven (IIHFWVRI(OHYDWHG=LQF&RQFHQWUDWLRQVLQ6RLO RQWKH3RWHQWLDO1LWULILFDWLRQ5DWH,GHQWLILFDWLRQRI%LRORJLFDO$YDLODELOLW\ DQG0HFKDQLVPVRI7ROHUDQFH Proefschrift voorgedragen tot het behalen van de graad van Doctor in de Bio-ingenieurswetenschappen door -HOOH0(57(16 0$$57

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3 ISBN Wettelijk depot D/2007/11.109/8

4 Katholieke Universiteit Leuven Faculteit Bio-ingenieurswetenschappen ',66(57$7,21(6'($*5,&8/785$ Doctoraatsproefschrift nr. 739 aan de faculteit Bio-ingenieurswetenschappen van de K.U.Leuven (IIHFWVRI(OHYDWHG=LQF&RQFHQWUDWLRQVLQ6RLO RQWKH3RWHQWLDO1LWULILFDWLRQ5DWH,GHQWLILFDWLRQRI%LRORJLFDO$YDLODELOLW\ DQG0HFKDQLVPVRI7ROHUDQFH 3URPRWRU Prof. E. Smolders, K.U.Leuven Prof. D. Springael, K.U.Leuven /HGHQYDQGHH[DPHQFRPPLVVLH Prof. G. Volckaert, voorzitter Prof. E. Bååth, Lund University Prof. J. Michiels, K.U.Leuven Prof. R. Merckx, K.U.Leuven Proefschrift voorgedragen tot het behalen van de graad van Doctor in de Bio-ingenieurswetenschappen door -HOOH0(57(16 0

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10 7DEOHRIFRQWHQWV 6DPHQYDWWLQJ 6XPPDU\ /LVWRIDEEUHYLDWLRQV &KDSWHU General introduction, hypotheses, general objectives and thesis outline 1 &KDSWHU Long-term exposure to elevated zinc concentrations induced structural changes and zinc tolerance of the nitrifying community in soil 13 &KDSWHU Zinc toxicity to nitrification in soil and soil-less solutions can be predicted with the same Biotic Ligand Model 33 &KDSWHU Recovery of the soil nitrification potential despite decreased diversity of the nitrifying community within 2 years after ZnSO 4 contamination in the field 51 &KDSWHU Resistance and resilience of zinc tolerant nitrifying communities to additional stressors is unaffected in long-term zinc contaminated soils 69 &KDSWHU General conclusions 77 5HIHUHQFHV 83 /LVWRISXEOLFDWLRQV 91 i v ix

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13 6DPHQYDWWLQJ Bodem microbiële processen zijn gevoelig aan verhoogde concentraties zware metalen, waaronder zink (Zn). Toevoegen van Zn 2+ zouten aan onverontreinigde bodems in het labo veroorzaakt toxische effecten bij microbiële processen, vaak al bij Zn concentraties die binnen de range van natuurlijke achtergrondwaarden liggen. Het is daarom absoluut noodzakelijk om kritische zinkconcentraties te bepalen waarbij het bodem ecosysteem permanente schade ondervindt. Maar in langdurig verontreinigde bodems worden er dikwijls pas toxische effecten gevonden bij zinkconcentraties die veel hoger zijn dan de toxische concentraties uit labotesten. In dit onderzoek wordt het verschil in Zn toxiciteit bij de nitrificatie in labo of langdurig gecontamineerde bodems bestudeerd. De hypothese is dat dit verschil te wijten is aan een verschillende biobeschikbaarheid van Zn in beide systemen, en aan microbiële adaptatie. Langdurig Zn verontreinigde bodems werden bemonsterd in weiland nabij een gegalvaniseerde elektriciteitspyloon in Zeveren (Gent). De onderliggende bodem werd aangerijkt met Zn door corrosie van de pyloon, en de concentraties stegen van 1 mmol Zn kg -1 in ongecontamineerde bodem tot 25 mmol Zn kg -1 in de meest gecontamineerde bodem onder de pyloon. De potentiële nitrificatiesnelheid (PNS) in deze bodems was niet beïnvloed door de verhoogde Zn concentraties, maar het toevoegen van Zn 2+ zout aan de ongecontamineerde bodem tot gelijke totale Zn concentraties veroorzaakte toxische effecten bij de PNS vanaf 1.8 mmol Zn kg -1. Het verschil in PNS tussen labo en langdurig gecontamineerde bodems wordt deels veroorzaakt door een verschillende Zn concentratie in poriewater. Zink tolerantie van de nitrificerende gemeenschap in langdurig verontreinigde bodem werd aangetoond door i) een lagere gevoeligheid van deze gemeenschap aan toegevoegd Zn dan de nitrificerende gemeenschap uit ongecontamineerde bodem na inoculatie van deze gemeenschappen in steriele, Zn verontreinigde bodem en doordat ii) de PNS van deze gemeenschap niet wijzigde na extra ZnCl 2 toevoeging, terwijl de PNS van de gemeenschap uit ongecontamineerde

14 ii 6DPHQYDWWLQJ bodem met 70% werd gereduceerd bij identieke Zn concentraties in oplossing. Bovendien werd met Denaturerende Gradiënt Gel Elektroforese (DGGE) aangetoond dat de structuur van de nitrificerende gemeenschap in de langdurig verontreinigde bodems veranderde met toenemende Zn concentraties. Er werd dus aangetoond dat de nitrificerende gemeenschap tolerant werd aan Zn door de langdurige blootstelling aan verhoogde Zn concentraties, en dat dit gepaard ging met een wijziging van de gemeenschapsstructuur. De afwezigheid van Zn toxiciteit in langdurig verontreinigde bodems werd deels verklaard door de verlaagde Zn concentraties in poriewater. Dit suggereert dat Zn biobeschikbaarheid gerelateerd is met de vrije Zn 2+ concentratie (of Zn 2+ activiteit) in oplossing. Deze hypothese werd getest door de effecten van ph en ionische samenstelling op Zn toxiciteit te vergelijken in een reeks labo verontreinigde bodems en in bodemloze culturen van 1LWURVRVSLUD sp. NpAV. De Zn concentratie in oplossing waarbij de PNS werd gereduceerd met 20% (Zn EC20 ) varieerde 150-voudig in bodem en 50-voudig in de bodemloze culturen. Bovendien steeg de Zn EC20 met toenemende H +, Ca 2+ en Mg 2+ concentraties in oplossing, wat een beschermend effect van deze kationen tegenover Zn toxiciteit suggereert. Zink speciatie werd bepaald in de bodems en de bodemloze culturen, en de vrije Zn 2+ activiteit (Zn 2+ ) EC20 bij de Zn EC20 werd gemodelleerd met een Biotisch Ligand Model (BLM) en een Freundlich-type model. Beide modellen houden rekening met de competitie van kationen voor de binding aan biotische liganden. De (Zn 2+ ) EC20 in bodem en bodemloze culturen werd voorspeld door één set parameters voor elk model. Beide modellen waren geschikt voor de modellering en voorspelden de geobserveerde (Zn 2+ ) EC20 binnen een factor 4. Deze resultaten suggereren dat de nitrificerende gemeenschap in bodem blootgesteld is aan Zn via het vrije Zn 2+ ion in oplossing. Eerder toonden we aan dat Zn tolerantie geïnduceerd werd door de langdurige blootstelling van de nitrificerende gemeenschap aan verhoogde Zn concentraties, maar het proces van adaptatie aan verhoogde Zn concentraties kon echter niet bepaald worden. Daarom werden bodems, gecontamineerd met toenemende ZnSO 4 concentraties, bemonsterd op verschillende tijdstippen in een veldexperiment in

15 6DPHQYDWWLQJ iii Spalding (Australië): binnen een week na toediening van de Zn zouten, en na 1, 2 en 3 jaar. Onmiddellijk na toedienen van de Zn zouten aan de bodem was de Zn concentratie waarop de PNS gereduceerd werd met 50% (Zn EC50 ) 15.5 mmol Zn kg -1, of 2.7 mm Zn in een 0.01 M CaCl 2 bodem extract. Deze waarden stegen tot 39.5 mmol Zn kg -1 en 9.3 mm in een 0.01 M CaCl 2 bodem extract na 2 jaar. Dit suggereert dat de nitrificerende gemeenschap in de Zn verontreinigde bodems adapteerde aan de verhoogde Zn concentraties binnen 2 jaar na Zn verontreiniging. Deze hypothese werd getest door de nitrificerende gemeenschap uit het ongecontamineerde bodemstaal en deze uit een Zn gecontamineerd bodemstaal (28.3 mmol toegevoegd Zn kg -1 ) van de verschillende staalnames bloot te stellen aan toenemende Zn concentraties in oplossing. Er werd duidelijk aangetoond dat het herstel van de PNS na 2 jaar gepaard ging met een verlaagde gevoeligheid van de nitrificerende gemeenschap in het Zn verontreinigd bodemstaal aan toenemende Zn concentraties. Het analyseren van de structuur van de nitrificerende gemeenschap met DGGE toonde een veranderde structuur en een afnemende diversiteit in de Zn verontreinigde bodemstalen in vergelijking met het ongecontamineerde staal. Er werd besloten dat het herstel van de PNS gepaard ging met Zn tolerantie, maar dat het niet gekoppeld was met een herstel van de diversiteit van de nitrificerende gemeenschap. De structuur van de nitrificerende gemeenschap veranderde met toenemende Zn concentraties in de langdurig gecontamineerde bodems uit het Zeveren transect, en de gemeenschap werd meer tolerant aan Zn. Het is echter onduidelijk of deze structurele wijziging en Zn tolerantie de gevoeligheid van het nitrificatie proces aan andere stress factoren beïnvloeden. Deze hypothese werd getest door de ongecontamineerde bodem en 4 langdurig gecontamineerde bodemstalen met toenemende Zn concentraties (4-24 mmol Zn kg -1 ) en dus ook toenemende Zn tolerantie, bloot te stellen aan additionele stressfactoren (een toxisch pesticide, vries-dooi cycli of droog-nat cycli). De weerstand van de nitrificerende gemeenschap aan de extra stress factoren werd bepaald door de PNS te meten onmiddellijk na blootstelling, en het herstelvermogen werd bepaald door de PNS te meten na 3 weken incubatie ( veerkracht ). De resultaten toonden aan dat zowel de weerstand als de veerkracht van de nitrificerende

16 iv 6DPHQYDWWLQJ gemeenschap aan de extra stress factoren niet werden aangetast door de langdurige blootstelling aan de verhoogde Zn concentraties. We toonden aan dat de nitrificerende gemeenschap in langdurig Zn verontreinigde bodems zich aanpaste aan de verhoogde Zn concentraties. Dit uitte zich in een verlaagde gevoeligheid aan Zn en een gewijzigde gemeenschapsstructuur. Bovendien werd er aangetoond dat nitrificerende microorganismen in de bodem blootgesteld werden aan Zn via vrije Zn 2+ ionen in oplossing. De combinatie van een verlaagde biobeschikbare Zn concentratie en een aanpassing van de nitrificerende gemeenschap aan verhoogde Zn concentraties is verantwoordelijk voor de afwezigheid van Zn toxiciteit in langdurig Zn verontreinigde bodems. Dit heeft echter geen gevolgen voor de gevoeligheid van het nitrificatie proces aan extra stress factoren.

17 6XPPDU\ Soil microbial processes are sensitive to elevated trace metal concentrations such as zinc (Zn). Soils that are amended with Zn 2+ salts in the laboratory reveal toxic effects at total soil Zn concentrations that are within the range of naturally occurring background Zn concentrations. These toxic effects are usually not identified in long-term Zn contaminated soil samples, even at total Zn concentrations that are several times larger than thresholds found in laboratory amended samples. The goal of this study is to identify the discrepancy of Zn toxicity to soil microbial processes between freshly amended soils and long-term contaminated soil samples. It is speculated that the discrepancy is related to differences in Zn bioavailability and to microbial adaptation. Long-term Zn contaminated soil was sampled near a galvanized electricity transmission tower in Zeveren (Belgium). Corrosion of the tower enriched the underlying soil with Zn, and total Zn concentrations increased from 1 mmol Zn kg -1 in uncontaminated soil to 25 mmol Zn kg -1 in the most contaminated soil under the tower. The Potential Nitrification Rate (PNR) in these soil samples was unaffected by the elevated Zn concentrations, whereas ZnCl 2 spiking of the uncontaminated soil to similar total Zn concentrations reduced the PNR significantly at 1.8 mmol Zn kg -1. The PNR in Zn spiked and long-term contaminated soil samples was partly explained by different Zn concentrations in the pore water. Zinc tolerance of the ammonia oxidizing bacteria (AOB) community in long-term contaminated soil was demonstrated by showing that i) the community was less sensitive to Zn than the AOB community of uncontaminated soil when they were inoculated in sterile ZnCl 2 spiked soil and ii) the PNR of the community was less affected by further ZnCl 2 additions whereas the PNR of the AOB community of uncontaminated soil was reduced by more than 70% at equivalent soluble Zn concentrations. Furthermore, DGGE fingerprinting showed that the AOB community structure in long-term contaminated soil changed with increasing Zn

18 vi 6XPPDU\ concentrations compared to the AOB community of uncontaminated soil. This field survey illustrated a so-called Pollution Induced Community Tolerance (PICT) which was associated with a change in community structure. Different Zn concentrations in pore water between long-term and freshly spiked soil at similar total Zn concentrations partly explained the different PNR in both systems with increasing total Zn concentrations. This suggests that Zn bioavailability is related to the Zn concentration (or Zn 2+ activity) in pore water. The hypothesis that soil solution Zn is the exposure pathway for soil micro-organisms was tested by comparing the effects of ph and ionic composition to Zn toxicity between a range of soils and soil-less solutions of 1LWURVRVSLUD sp. NpAV. The Zn concentration at which the nitrification was reduced by 20% (Zn EC20 ) varied 150-fold between soils and 50-fold between soil-less solutions, and decreased with increasing H +, Ca 2+ and Mg 2+ concentrations in both systems, suggesting protective effects of these cations against Zn toxicity. Zn speciation was determined, and the Zn 2+ activities at the Zn EC20 ((Zn 2+ ) EC20 ) in both systems were modelled using a Biotic Ligand Model and a Freundlich-type model. Both models take into account cation competition for Zn toxicity. The (Zn 2+ ) EC20 of soil and soil-less solutions were predicted using a single set of parameters per model. Both models performed equally well, and observed (Zn 2+ ) EC20 were predicted within a factor 4. These results suggest that nitrifying organisms are exposed to Zn through the free Zn 2+ ion in solution. The field survey showed that long-term exposure of the nitrifying community to elevated Zn concentrations increased its Zn tolerance. No information was obtained about the adaptation rate of the AOB community. Therefore, soils with increasing ZnSO 4 amendments were sampled in a field trial in Spalding (Australia) within one week after metal addition and after 1, 2 and 3 years. Initially, the PNR was reduced by 50% (Zn EC50 ) at 15.5 mmol Zn kg -1 or at 2.7 mm Zn in a 0.01 M CaCl 2 soil extract. The Zn EC50 increased to 39.5 mmol Zn kg -1 and 9.3 mm Zn after two years incubation in the field, suggesting Zn tolerance of the AOB community within 2 years after Zn spiking. This hypothesis was tested by adding different amounts of ZnCl 2 to suspensions of uncontaminated and Zn contaminated (28.3 mmol added Zn kg -1 ) soil of the different

19 6XPPDU\ vii sampling periods (i.e. soil that was sampled within one week after Zn spiking and after 1, 2 and 3 years incubation in the field). We showed that recovery of the PNR within 2 years after Zn addition was coupled to a decreased sensitivity of the AOB community to increasing Zn concentrations. DGGE fingerprinting revealed a changed AOB community structure and a decreased AOB diversity in Zn spiked, aged soil compared to uncontaminated soil. It was concluded that recovery of the nitrification was related to the development of Zn tolerance which was associated with a change of the AOB community. The development of tolerance (PICT) clearly indicates an effect of exposure to a toxic substance, but its ecological consequences are unclear. The relationship between Zn tolerance and functional stability after applying additional stress factors was, therefore, further explored. The uncontaminated soil sample and 4 long-term Zn contaminated soil samples with increasing Zn concentrations (4-24 mmol Zn kg -1 ) and, hence, increasing Zn tolerance were exposed to a toxic pesticide (metamitron) or to three freeze-thaw or three dry-wet cycles. The immediate response of the soil samples ( resistance ) and the ability of the soil samples to recover from the stressors after 3 weeks incubation ( resistance ) were measured. It was found that, in general, tolerance did not negatively affect the functional stability of the AOB community. Overall, it is concluded that the AOB community in long-term Zn contaminated soil adapts to elevated Zn concentrations, as shown by i) a reduced Zn sensitivity and ii) a changed community structure, and that the AOB community is exposed to Zn through the free Zn 2+ ion in solution. The combination of Zn adaptation and a decreased Zn bioavailability reduces the toxic effect of Zn in long-term Zn contaminated soil samples compared to the effects of Zn in freshly spiked soil samples. Adaptation of the AOB community to elevated Zn concentrations does not influence its sensitivity towards additional stressors.

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21 /LVWRIDEEUHYLDWLRQV AGE AOB BLM CEC C total DAPI DGGE DOC DOM DNA dntp EC X FIAM HC HEPES HPLC HPVC Agarose Gel Electrophoresis Ammonia Oxidizing Bacteria Biotic Ligand Model Cation Exchange Capacity Total carbon concentration 4,6-diamidino-2-phenylindole dihydrochloride Dissolved Organic Carbon Dissolved Organic Matter Deoxyribonucleic Acid Deoxyribonucleoside Triphosphate Effective Concentration causing x% reduction: contaminant concentration at which the response is reduced by x% compared to the response without exposure to the contaminant Free Ion Activity Model Hazardous Concentration Denaturing Gradient Gel Electrophoresis 4-(2-hydroxyethyl)piperazine-1- ethanesulfonic acid High Performance Liquid Chromatography High Production Volume Chemical ICP-OES Inductively Coupled Plasma- Optical Emission Spectroscopy LC X Lethal Concentration causing x% lethality: contaminant concentration at which x% lethality occurs MES 2-(N-morpholino)ethanesulfonic acid MOPS MPN MWHC NIST NOEC N total OTU PCA PCR PICT PLFA PNEC PNR RFLP RNA SIN STP tblm TIT WHAM Zn ECX Zn pw Zn total 3-(N-morpholino)propanesulfonic acid Most Probable Number Maximum Water Holding Capacity National Institute of Standards and Technology No Observed Effect Concentration: highest contaminant concentration at which the response is not significantly affected Total nitrogen concentration Operational Taxonomic Unit Principal Component Analysis Polymerase Chain Reaction Pollution Induced Community Tolerance Phospholipid Fatty Acid Potential No Effect Concentration Potential Nitrification Rate Restriction Fragment Length Polymorphism Ribonucleic Acid Substrate Induced Nitrification Sewage Treatment Plant Terrestrial Biotic Ligand Model Thymidine Incorporation Technique Windermere Humic Acid Model Zinc concentration at the EC X Zinc concentration in pore water Total zinc concentration

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23 &+$37(5 *HQHUDOLQWURGXFWLRQK\SRWKHVHVJHQHUDOREMHFWLYHVDQGWKHVLVRXWOLQH This thesis addresses the effects of zinc (Zn) to a soil microbial process, i.e. soil nitrification. It explores the differences in toxicity response between short- and longterm exposure in terms of Zn bioavailability and the development of Zn tolerant communities by adaptation. Finally, the ecological effects of the adaptation process are assessed. Toxicity of Zn to soil microbial processes is one of the most critical factors in Zn contaminated soils, i.e. risks of Zn in soil might be triggered by these processes. This chapter summarises the Zn concentrations in soil, their enrichment by anthropogenic emissions, the toxic effects of Zn to soil microbial processes and the role of microbial adaptation on the ecotoxicological effects of Zn on the long term. 7RWDO]LQFFRQFHQWUDWLRQLQVRLOV Zinc can be present in different minerals of the parent rock of a soil. Zinc containing minerals include sphalerite, smithsonite, zinc spar and marmatite. Zinc can also be present as a trace constituent of different other minerals. Soil formation is associated with weathering of the parent soil material. Zinc released from weathering rocks is adsorbed by soil organic matter, by oxyhydroxides or can precipitate in secondary minerals. The naturally occurring Zn concentration in soil ( geological background concentration ) therefore depends on the mineral composition of the parent material. It is generally found that soil background concentrations of trace metals are positively correlated with the clay and organic matter content. Soils are, however, also historically enriched by Zn from agricultural practises or industrial activities. Therefore, current background Zn concentrations in soils well away from point sources ( ambient concentrations ) exceed the geological background concentrations (Reimann and Garrett, 2005). Consequently, the ambient Zn concentrations are both affected by the

24 2 &KDSWHU geological properties of the soil and by the land use history. Selected Zn concentration ranges are shown in Table 1. 7DEOH Mean and range of total Zn concentrations (mg kg -1 ) in unpolluted soils of European countries. %HOJLXP 7KH1HWKHUODQGV 'HQPDUN $XVWULD )UDQFH Mean Range De Temmerman et al. (2000); 2 Angelone and Bini (1992); 3 Baize (1997),QGXVWULDODSSOLFDWLRQRI=QDQGUHOHDVHWRWKHHQYLURQPHQW Zinc is an economically important metal used in various applications. Its production in the European Union (EU) was about 2.1*10 6 Mg per year in (EU Risk Assessment Report Part I: Environment). In 1997, more than 35% was applied for galvanizing metal constructions to protect them against oxidation. Furthermore, about 25% was used for brass production (brass contains 20-45% Zn), 10% for alloy and die casting, 10% for the production of rolled and wrought Zn and 3% for the production of Zn powder. The remaining 10% was used for the production of Zn compounds as e.g. ZnO for paints, rubber, cosmetics and ZnS for luminous dials, paints, fluorescent lights. The production and consumption of Zn or Zn based materials release Zn to the environment. The major sources of environmental anthropogenic emissions are corrosion of galvanised structures, application of manure (Zn is added as micronutrient in the animal feed) or sewage sludge, tyre wear and oil and coal combustion (Cleven et al., 1993). Industrialisation of the Western world severely increased the SO 2 concentrations in air, resulting in a significant contribution of atmospheric corrosion to the total Zn emissions to water and soil. Zinc emissions from corrosion were about 30% (73 Mg y -1 ) of the total Zn emissions to surface water and 1% of the total Zn emissions to soil in the Netherlands in 1999 (EU Risk Assessment Report Part I: Environment). An overview of the Zn emissions to water, soil and air in the Netherlands in 1999 is shown in Table 2. The net input of Zn in soil is mainly derived from manure application. De Vries et al. (2004) reported diffuse Zn emissions to agricultural soils in

25 &KDSWHU 3 the Netherlands of g Zn ha -1 y -1 which increase total soil Zn in the topsoil by about 0.5 mg Zn kg -1 annually. Locally higher emissions may obviously occur, e.g. in the soils receiving sludge and soils under galvanised constructions (e.g. electricity transmission towers). Lijzen and Franken (1994) reported Zn concentrations of mg Zn kg -1 in the subsoil under electricity transmission towers in the Netherlands, whereas Zn concentrations increased to mg kg -1 under pylons in the U.K. and Canada. The Zn concentrations depended on the level of air pollution (SO 2 concentration), climate (rainfall) and the age of the pylon (duration of emission). Historical point emissions of Zn smelters have resulted in elevated soil Zn concentrations in these areas such as in the Kempen area of Belgium and The Netherlands. 7DEOH Zinc emissions to water, soil and air in the Netherlands (Mg y -1 ) in 1999 (De Vries et al., 2004). :DVWHZDWHU 6XUIDFH 6RLO ZDWHU Agriculture $LU Industry Waste treatment 4 Traffic Consumers Trade and services (emissions of educational institutes, recreation, ) 37 2 Effluents sewage treatment plants (STP) 95 Others (use of sewage sludge, ) Atmospheric deposition RWDO 7KHUROHRI]LQFLQELRWD The essentiality of Zn has been demonstrated for all living organisms. In prokaryotes, Zn is essential as cofactor in enzymes (e.g. carbonic anhydrase, alcohol dehydrogenase, RNA and DNA polymerase) and for the maintenance of proteine structure (Nies, 1999; Blencowe and Morby, 2003). The Zn uptake and efflux mechanisms are strictly controlled by regulators at the level of transcription, modulating

26 4 &KDSWHU gene expression in response to the internal Zn concentration, to maintain the cytoplasmic Zn concentration within narrow limits independenty of the external Zn concentration. Zinc uptake mechanisms may be either slow and Zn specific, coupled to an energy source (e.g. ZnuABC transporters), or fast and non-specific, driven by a concentration gradient over the membrane (e.g. CorA proteins). Export systems are mostly proteins that transport metal ions across the cytoplasmic membrane, and can rely on the use of chemical energy or the energy, stored in electrochemical gradients e.g. CadA, a P-type ATPase, or the czc-system, a chemiosmotic transenvelop transporter (Blencowe and Morby, 2003; Nies, 2003). Soil contamination by Zn increases the Zn concentration in the soil solution. Unspecific transporters are constitutively expressed in prokaryotic cells, and this gate can not be closed. The rapid influx of Zn ions increases the optimal Zn concentration and once the optimal Zn concentration of a microorganism is exceeded, the beneficial effects of Zn disappear and Zn becomes toxic (Nies and Silver, 1995; Figure 1). Toxic effects of Zn are mostly related to the displacement of essential metals in enzymes (e.g. Zn 2+ for Mg 2+ ), excessive binding to SH groups and inactivating the enzyme or binding to glutathione in Gram negative bacteria, resulting in extra oxidative stress (Nies, 1999). The cell can respond to the excess of cytoplasmic Zn ions by diminishing the expression of genes for fast and non-specific Zn import or by the transcription of metal-responsive regulators that induce the expression of genes for Zn export mechanisms. Alternatively, Zn can be intracellularly of extracellularly sequestered to metal binding proteins (e.g. metallothionein), reducing the exposure of the cell to Zn 2+ ions. Some of these resistance mechanisms are chromosome-based, while other can be plasmid-encoded (Nies, 1999; Bruins et al., 2000). Yet, no general resistance mechanism for all metals exists (Silver, 1998). The essentiality and toxicity of Zn result in a dose-response curve with a typical pattern as given in Figure 1. Various soils are Zn deficient worldwide as assessed by plant responses to Zn fertilisation. Soil parameters such as increased soil ph, CaCO 3 or organic matter content decrease the phytoavailability of Zn, resulting in reduced yields, impaired quality of plants or chlorosis on millions of hectares of arable land worldwide.

27 /!, &' "! &KDSWHU 5 This may not only affect crop production, but also human Zn uptake. Important food crops (rice, maize and sorghum) are very susceptible to Zn deficiency, and may severely reduce human Zn provision (± mg Zn day -1 ) in developing countries. More than one third of the human population (>2*10 9 ) is potentially affected by Zn deficiency, causing e.g. growth retardation, delayed wound healing or diarrhoea (Alloway, 2003). In contrast, elevated Zn concentrations in soil result in agricultural products with unacceptable Zn levels violating food quality criteria, and toxic effects towards earthworms, nematodes and microorganisms (De Vries et al. 2004). The effects of elevated Zn concentrations towards soil microorganism are further explored in this thesis. ((&$ *! #$.) *& -#! +#$ (#$ *#) %& #$ )LJXUH The effect of increasing metal concentration in the environment to organisms for essential and inessential metals (derived from Gadd, 1992). 7R[LFHIIHFWVRI=QIRUVRLORUJDQLVPV Elevated soil Zn concentrations can exert toxic effects towards soil microorganisms, plants and soil invertebrates. The risk of Zn to higher organisms and humans via food chain transfer is relatively small because plants and invertebrates are more readily affected by elevated soil Zn and, hence, limited food is produced in soils with elevated Zn. Therefore, critical Zn concentrations in soils are triggered by ecotoxicological

28 6 &KDSWHU effects on soil organisms. Various attempts have been made to determine soil Zn concentrations at which the soil ecosystem is protected against adverse effects (Bååth, 1989; Giller et al., 1998). In 1993, the Council Regulation 793/93/EEC (EC, 1993) was introduced by the European community to ensure the save use of existing chemicals (i.e. chemicals that were on the market in 1981). Initially, the EC regulation was only focussed on high production volume chemicals (HPVCs; production > 1000 Mg y -1 ), and Zn (metallic Zn, Zn oxide, Zn chloride, Zn stearate, Zn sulphate, Zn phosphate) was the first metal that was added to the second priority list in A major aim of the risk assessment was the determination of toxic Zn thresholds for soil organisms by reviewing the existing literature and by conducting additional experiments. Data were collected from studies in which Zn salt was added to uncontaminated soil at different doses and a compilation was made of the highest added Zn concentrations at which no significant toxic effects towards the soil organisms were observed (No Observed Effect Concentration or NOEC ). These values are relatively smaller for soil microbial processes than for plants or invertebrates. The NOEC values for some microbe mediated processes in soil that were used for the Zn risk assessment are shown in Table 3. The Predicted No Effect Concentration or PNEC was calculated by plotting the NOECs of the microbe mediated processes, expressed as added Zn concentrations, in a cumulative probability curve and deriving the 5 th percentile of the curve (Figure 2; HC 5 ). A critical value of 27 mg added Zn kg -1 soil was obtained. The concept of the PNEC, derived by this so-called species sensitivity distribution, is that 95% of all soil microorganisms or soil microbial processes are protected at that Zn concentration (Aldenberg and Slob, 1993). In practice, a soil is considered to be potentially affected by Zn when the total soil Zn added to a soil exceeds 27 mg Zn kg -1, i.e. in soils with ambient Zn concentration 27 mg Zn kg -1 above background. From Table 1 it is clear that this PNEC is within the range of background concentration and it is practically impossible to discriminate added Zn from geogenic Zn in such a small concentration range. This example illustrates the

29 &KDSWHU 7 relatively large sensitivity of soil microbial processes when relying on studies with Zn salt spiked soils (Bodar et al., 2005). 7DEOH Range of NOEC values, expressed as added Zn concentrations, for microbial and enzymatic soil processes. Process Range (mg Zn kg -1 ) Reference C mineralization Cornfield, 1977 Lighthart et al., 1983 N mineralization Chang and Broadbent, 1982 Necker and Kunze, 1986 Nitrification Smolders et al., 2003 Mineralization of glucose Smolders et al., 2003 Mineralization of glutamic acid Notenboom and Posthuma, 1995 Haanstra and Doelman, 1984 Mineralization of maize residue 50 - >1800 Smolders et al., 2003 Arylsulphatase activity Haanstra and Doelman, 1991 Dehydrogenase activity Maliszewska et al., 1985 Phosphatase activity Doelman and Haanstra, 1985 Urease activity Doelman and Haanstra, 1986 &XPXODWLYH SUREDELOLW\ +& 78 ORJ12(&PJNJ )LJXUH Cumulative probability plot of the NOEC concentrations for added Zn (log transformed), based on toxicity data of soil microbial mediated processes with indication of the HC 5 value (from: EU Risk Assessment Report Part I: Environment).

30 B A B 8 &KDSWHU Field data, however, reveal that toxic Zn concentrations are well above those obtained when spiking corresponding soils with Zn 2+ salts to identical total Zn concentrations. (Smolders et al., 2003; Smolders et al., 2004). For example, no toxic effects on the nitrification were observed in field soils containing more than 3000 mg Zn kg -1, whereas spiking an uncontaminated soil of the same site with ZnCl 2 reduced the nitrification already at 300 mg added Zn kg -1 (Smolders et al., 2004; Figure 3). This discrepancy between spiked soil samples and long-term contaminated soil samples is attributed to i) a different Zn bioavailability in both systems and ii) adaptation of the soil microorganisms to the elevated Zn concentrations. G NJ 315PJ12 =Q92:;9 <= PJNJ> )LJXUH Potential Nitrification Rate (PNR) in the long-term contaminated soil samples (closed symbols) and the ZnCl 2 spiked soil samples (open symbols) with increasing total Zn concentration (from Smolders et al., 2004). =LQF WR[LFLW\ WR VRLO PLFURELDO SURFHVVHV DV DIIHFWHG E\ VRLO SURSHUWLHV DQG HTXLOLEUDWLRQWLPHDIWHU]LQFFRQWDPLQDWLRQ The data in Table 3 and Figure 3 indicate that Zn toxicity data are highly variable when assessed with different microbial assays in different soils. This observation suggests that the inhibitory effects of Zn are not uniform in all soils and depend on soil properties and on the equilibration time after spiking. Added Zn will distribute over the solid and liquid soil phase, and Zn can be present as free Zn 2+ ions in the soil solution or

31 &KDSWHU 9 bound to organic or inorganic ligands or to solid soil particles. The distribution of Zn over the two compartments is quantified by the. d coefficient, representing the ratio of the Zn concentration on the solid phase over the amount of Zn in the liquid phase (i.e. the soil solution). The. d coefficient is mainly determined by the soil ph, as the Zn concentration in solution increases 2-8 fold as the soil ph decreases by 1 unit (Sauvé et al., 2000; Degryse et al., 2003). It is generally stated that the soil microflora is exposed to metals via the pore water and the free metal ion is more available than the metal complexes, however, a direct proof has not yet been given (Janssen et al., 1997). Therefore, soil properties that control the fraction of total soil Zn that is present as Zn 2+ in solution might affect the relationship between the total Zn concentration and the effects. Ageing processes increase the amount of metal bound to the solid phase on the long-term. As a result, the metal concentration in solution, and thus the bioavailable and potentially toxic concentration, decreases over time. This can be a major factor to explain the lack of Zn toxicity in long-term contaminated soil samples, compared to freshly contaminated soil samples with similar total Zn concentrations (Giller et al., 1998). Microorganisms in uncontaminated soil samples maintain a constant internal Zn concentration through equilibrium between Zn influx and Zn efflux mechanisms. As discussed above, background Zn concentrations in uncontaminated soil samples vary, but soil microorganisms are adapted to these varying Zn concentrations (McLaughlin and Smolders, 2001). Moreover, these authors showed a positive relationship between soil background Zn concentrations and the NOEC values of Zn toxicity studies for a range of microbial and enzymatic assays. This suggests that the sensitivity of the soil microbial community to elevated Zn concentrations depends on its previous exposure to Zn. This confirmed previous observations by e.g. Bååth et al. (1992b) or Diáz-Raviña et al. (1994) that long-time exposure of the soil microbial community to elevated Zn concentrations in Zn spiked soils increases its resistance to extra Zn additions. The adaptation process to artificially increased Zn concentrations likely occurs through extinction of Zn sensitive species that are immediately killed (e.g. membrane disruption) or that are not able to change their Zn metabolism accurately on the longterm (e.g. blockage of critical cellular functions). Inherently Zn tolerant species or

32 10 &KDSWHU species that changed their Zn metabolism sufficiently to maintain a normal cytoplasmic Zn concentration will proliferate and colonize all available niches. These resistance mechanisms, as well as the time needed to induce them, can differ significantly between species (Diáz-Raviña and Bååth, 1996; Bruins et al., 2000). The discrepancy between Zn spiked and long-term contaminated soil samples is relatively large (Figure 3) and there is a dire need to explain that difference to better define the environmental risk of Zn. It could be advocated that field data should be the basis to set soil limits, however the inherent presence of confounding factors in field contaminated soils (mixed contamination, correlation between metal contamination and soil organic matter concentrations) obscure cause-effect relationships. Therefore, this work will assess the mechanisms by which effects on long-term contaminated soils differ from that in freshly spiked soils in terms of Zn bioavailability and adaptation of the soil microbial community to the increased Zn concentrations. In addition, the role of Zn speciation in soils on the Zn toxicity will be assessed to validate the concept of pore water exposure to Zn 2+ as the toxic route for microorganisms. We focus on the nitrification process in soil, because it is a specific metabolic process of the N-cycle that is supposed to be carried out by a discrete microbial community, that can be easily measured (Koops et al., 2003) and that is sensitive to heavy metal stress (Smolders et al., 2001; Broos et al., 2005). Recently, Leininger et al. (2006) quantified the copy number of the bacterial and the archaeal ammonia monooxygenase (DPR$) gene in 12 uncontaminated soil samples, and observed a higher abundance of archaeal DPR$ gene copies than bacterial DPR$ gene copies. This observation suggests that crenarchaeota may be the most abundant ammonia oxidizing organisms in uncontaminated soils. Yet, these authors did not elucidate the relative contributionsof crenarchaeota and bacteria to the nitrification process in these soils.

33 &KDSWHU 11 +\SRWKHVHVDQGJHQHUDOREMHFWLYHV The K\SRWKHVHV of this thesis are: (i) The immediate effects of Zn additions to soil are not representative for the effects on the long-term (Renella et al., 2002). (ii) Microbial communities are exposed to Zn through the free Zn 2+ ions in solution (Janssen et al., 1997) but cations in solution (including H + ions) might compete with Zn 2+ for Zn uptake and toxicity as in aquatic studies. Consequently, effects of ph and ionic composition on soil solution Zn 2+ thresholds to microbial processes are similar between soils and soil-less cultures. (iii) Microbial communities can adapt to elevated Zn concentrations. As a result, its sensitivity towards Zn is reduced after prolonged exposure and the microbial process will be restored to the level in the undisturbed soil (Bååth et al., 1992b; Diáz-Raviña and Bååth, 1996). (iv) The adaptation process of microbial communities to Zn affects its ability to cope with additional disturbances or stress, as suggested by macro-ecological studies (e.g. Tillman, 1996). Based on these hypotheses, the main REMHFWLYHV are: (i) to explain the different responses of the nitrifying community between fresh Zn contaminations and long-term contaminations in the field at similar Zn contamination levels in terms of Zn bioavailability and adaptation of the nitrifying community. (ii) to create a model to predict Zn toxicity in Zn spiked soils and in Zn spiked soilless cultures to verify that nitrifying microorganisms are exposed to Zn through the free Zn 2+ ions in solution.

34 12 &KDSWHU (iii) to assess the rate of adaptation of a nitrifying community to elevated Zn in field conditions. (iv) to quantify the influence of Zn tolerance to the sensitivity of the nitrifying community to additional stress or disturbances. 7KHVLVRXWOLQH The thesis is divided in 4 chapters, each studying one of the 4 objectives. In &KDSWHU, the nitrification is measured in long-term field contaminated and in freshly ZnCl 2 spiked soil samples. The nitrification in both systems is related to total Zn concentrations and to Zn concentrations in pore water in both systems. Zinc tolerance of the nitrifying community is assessed by measuring the sensitivity of the original and the Zn exposed community towards increasing Zn concentrations, and by applying molecular techniques. In &KDSWHU, the nitrification is measured in a range of freshly ZnCl 2 contaminated soils, and in ZnCl 2 amended soil-less cultures of 1LWURVRVSLUD cells. Toxic concentrations are calculated, and a model is constructed to predict Zn toxicity in both systems based on the ph and the ionic composition of the solution. In &KDSWHU, the nitrification is monitored in ZnSO 4 spiked soil samples that are incubated under natural conditions for 3 years. Adaptation is assessed by monitoring the nitrification process in the Zn spiked soil samples over time and relating the activity of the nitrifying community to its Zn sensitivity and to the community structure. In &KDSWHU, the sensitivity of unexposed and Zn adapted nitrifying communities towards additional stressors (pesticide addition, freeze-thaw cycles or dry-wet cycles) is compared.

35 &+$37(5 /RQJWHUPH[SRVXUHWRHOHYDWHG]LQFFRQFHQWUDWLRQVLQGXFHGVWUXFWXUDO FKDQJHVDQG]LQFWROHUDQFHRIWKHQLWULI\LQJFRPPXQLW\LQVRLO $GDSWHGIURP0HUWHQV-6SULQJDHO''H7UR\HU,&KH\QV.:DWWLDX36PROGHUV (/RQJWHUPH[SRVXUHWRHOHYDWHG=QFRQFHQWUDWLRQVLQGXFHGVWUXFWXUDOFKDQJHVDQG =QWROHUDQFHRIWKHQLWULI\LQJFRPPXQLW\LQVRLO(QYLURQ0LFURELRO,QWURGXFWLRQ As mentioned in the Chapter 1, there is a large uncertainty about the derivation of critical thresholds for toxic trace metals to protect the soil microbial community. Toxicity data from field contaminated soil are often subject to interpretation due to mixed contamination and the covariance of the contamination level with other soil properties that also affect the soil microbial community. Alternatively, metals can be added to the soil in the laboratory and the immediate response (i.e. hours to weeks) of a microbial process to increasing metal concentrations can be assessed. It has repeatedly been found that there is a large discrepancy between the critical concentrations found by either method (Giller et al., 1998; Smolders et al., 2004). Nitrification has a key position in the nitrogen cycle and is one of the most sensitive soil microbial processes with regard to heavy metal stress (Koops et al., 2003; Broos et al., 2005). For example, the total Zn concentration at which the Potential Nitrification Rate (PNR) is reduced by 50% of its control value (=Zn EC50 ) in a freshly ZnCl 2 spiked soils is more than tenfold below the corresponding concentrations required to have a similar impact on C-mineralization endpoints (data of the same soil as used in this study here, Smolders et al., 2004). Moreover, Zn additions to reduce the nitrification rate by 10% is typically only 2-3 fold above the background concentration of the soil (Smolders et al., 2004). However, PNRs in long-term Zn contaminated soil samples were insensitive to total soil Zn concentrations that were 1-8 fold larger than corresponding Zn EC50 values of freshly spiked soils (Smolders et al., 2003, 2004). The discrepancy

36 14 &KDSWHU between metal toxicity in freshly spiked and long-term metal contaminated soils can be attributed to three factors: bioavailability, ionic strength and community adaptation. With regard to bioavailability of metals in freshly spiked and long-term metal contaminated soils, Zn concentrations in pore water of field contaminated soils were always less than 0.6 mm Zn while these concentrations were up to 6.9 mm Zn in soil spiked with ZnCl 2 to identical total Zn concentrations in soil (Smolders et al., 2003). Ionic strength increases after addition of metal salt solutions to soil and creates an osmotic shock to micro-organisms resulting in a decreased activity (Sindhu and Cornfield, 1967; Ping et al., 1996; McLaughlin et al., 2004). For example, increasing the electric conductivity of soil to 1 ds m -1 by adding NaCl or Na 2 SO 4 inhibited the nitrification (Inubushi et al., 1999). Changes of the microbial community structure in response to metal contamination have been previously shown using different methods for studying community properties (Díaz-Raviña et al., 1994; Frostegård et al., 1996; Kelly et al., 1999; Díaz-Raviña and Bååth, 2001; Van Beelen et al., 2001; Davis et al., 2004). However, most of these studies used methods rely on comparing the activity or the composition of the entire soil microbial community. Less is known about the adaptation of a well defined community, characterized by a specific metabolism, to metal contamination. Recently, Rusk et al. (2004) demonstrated the acquisition of Zn tolerance of the nitrifying community in soil spiked with ZnSO 4 up to 13.6 mmol Zn kg -1. After an incubation period of 17 months, the nitrifying community of the Zn contaminated soil was less sensitive to Zn than the nitrifying community of the uncontaminated control soil. Yet, the field relevance of this experiment is unclear because the Zn spiked soil was contaminated by a single addition of ZnSO 4 while in the field, metals are added progressively. In addition, it is unclear whether or not tolerance acquisition was associated with changes in nitrifying community structure. Stephen et al. (1999) and Gremion et al. (2004) contaminated soils with a mixture of toxic metals (respectively Cd, Co, Cs and Sr and Zn, Cu and Cd) and monitored the ammonia oxidizing bacteria (AOB) community structure. They both concluded that the AOB community changed due to long-term exposure to the metals (respectively 8 weeks and 12 months). Gremion et al. (2004) related this observation only to the potential ammonium oxidation of the AOB community while Stephen et al. (1999) did not

37 &KDSWHU 15 measure metal resistance or activity of the nitrifying community. So far, no studies have demonstrated metal tolerance of the nitrifying community in long-term metal contaminated soils or that metal tolerance is associated with shifts of the community structure. The objective of this chapter was to explain the large discrepancy of the nitrification response between freshly spiked soils and long-term contaminated soils in terms of Zn bioavailability, ionic strength in pore water and Zn tolerance of the nitrifying community. Therefore, nitrification rates in soils spiked with ZnCl 2 and CaCl 2 were compared with those in long-term contaminated soils and the results were related to Zn concentrations in pore water and ionic strength in pore water. In addition, the different Zn tolerance of the nitrifying community between long-term contaminated soils and reference soils was examined by measuring the response to a fresh Zn contamination. Finally, all observations were related to changes of the AOB community structure in response to the elevated Zn concentrations. 0DWHULDOVDQG0HWKRGV 6RLOVDPSOHV Topsoil (0-5 cm) was collected in grassland at different locations along a long-term Zn contaminated transect near a galvanized electricity transmission tower in Zeveren (Belgium). It was a sandy loam soil with a cation exchange capacity (CEC) of cmol c kg -1 (Smolders et al., 2004). The pylon was erected in 1992 and painted in As a result, the underlying soil was progressively enriched by corroded Zn for more than 10 years before sampling. Zn concentrations decreased with increasing distance from the pylon. The uncontaminated control soil containing background Zn concentrations was sampled 30 m away from the pylon. Soil sampling occurred in February 2003 (series 1) and October 2004 (series 2). Fresh soil was sieved immediately (< 4 mm) to remove stones and organic material and stored until use in plastic boxes at 4 C in the dark.